Subscriber access provided by UNIV OF CALIFORNIA SAN DIEGO LIBRARIES
Article
From Sediment to Top Predators: Broad Exposure of Polyhalogenated Carbazoles in San Francisco Bay (U.S.A.) Yan Wu, Hongli Tan, Rebecca Sutton, and Da Chen Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b05733 • Publication Date (Web): 23 Jan 2017 Downloaded from http://pubs.acs.org on January 23, 2017
Just Accepted “Just Accepted” manuscripts have been peer-reviewed and accepted for publication. They are posted online prior to technical editing, formatting for publication and author proofing. The American Chemical Society provides “Just Accepted” as a free service to the research community to expedite the dissemination of scientific material as soon as possible after acceptance. “Just Accepted” manuscripts appear in full in PDF format accompanied by an HTML abstract. “Just Accepted” manuscripts have been fully peer reviewed, but should not be considered the official version of record. They are accessible to all readers and citable by the Digital Object Identifier (DOI®). “Just Accepted” is an optional service offered to authors. Therefore, the “Just Accepted” Web site may not include all articles that will be published in the journal. After a manuscript is technically edited and formatted, it will be removed from the “Just Accepted” Web site and published as an ASAP article. Note that technical editing may introduce minor changes to the manuscript text and/or graphics which could affect content, and all legal disclaimers and ethical guidelines that apply to the journal pertain. ACS cannot be held responsible for errors or consequences arising from the use of information contained in these “Just Accepted” manuscripts.
Environmental Science & Technology is published by the American Chemical Society. 1155 Sixteenth Street N.W., Washington, DC 20036 Published by American Chemical Society. Copyright © American Chemical Society. However, no copyright claim is made to original U.S. Government works, or works produced by employees of any Commonwealth realm Crown government in the course of their duties.
Page 1 of 35
Environmental Science & Technology
1
From Sediment to Top Predators: Broad Exposure of Polyhalogenated
2
Carbazoles in San Francisco Bay (U.S.A.)
3 Yan Wu, † Hongli Tan, ‡,† Rebecca Sutton, § Da Chen ‡,†,*
4 5 6
†
7
Illinois University, Carbondale, Illinois 62901, United States
8
‡
9
Health, and Guangdong Key Laboratory of Environmental Pollution and Health, Jinan
Cooperative Wildlife Research Laboratory and Department of Zoology, Southern
School of Environment, Guangzhou Key Laboratory of Environmental Exposure and
10
University, Guangzhou, 510632, China
11
§
12
United States
San Francisco Estuary Institute, 4911 Central Avenue, Richmond, California 94804,
13 14 15 16 17 18 19 20 21 22
*Corresponding author email:
[email protected]; phone: (618) 453-6946; fax: (618) 453-
23
6944.
ACS Paragon Plus Environment
1
Environmental Science & Technology
Page 2 of 35
24
Abstract. The present study provides the first comprehensive investigation of
25
polyhalogenated carbazoles (PHCZ) contamination in an aquatic ecosystem. PHCZs have
26
been found in soil and aquatic sediment from several different regions, but knowledge of
27
their bioaccumulation and trophodynamics is extremely scarce. This work investigated a
28
suite of 11 PHCZ congeners in San Francisco Bay (United States) sediment and
29
organisms, including bivalves (n = 6 composites), sport fish (n = 12 composites), harbor
30
seal blubber (n = 18), and bird eggs (n = 8 composites). The most detectable congeners
31
included 3,6-dichlorocarbazole (36-CCZ), 3,6-dibromocarbazole (36-BCZ), 1,3,6-
32
tribromocarbazole (136-BCZ), 1,3,6,8-tetrabromocarbazole (1368-BCZ), and 1,8-
33
dibromo-3,6-dichlorocarbazole (18-B-36-CCZ). The median concentrations of ΣPHCZs
34
were 9.3 ng/g dry weight in sediment and ranged from 33.7 to 164 ng/g lipid weight in
35
various species. Biomagnification was observed from fish to harbor seal and was mainly
36
driven by chlorinated carbazoles, particularly 36-CCZ. Congener compositions of PHCZs
37
differed among species, suggesting that individual congeners may be subject to different
38
bioaccumulation or metabolism in species occupying various trophic levels in the studied
39
aquatic system. Toxic equivalent (TEQ) values of PHCZs were determined based on their
40
relative effect potencies (REP) compared to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD).
41
The median TEQ was 1.2 pg TEQ/g dry weight in sediment and 4.8 – 19.5 pg TEQ/g
42
lipid weight in biological tissues. Our study demonstrated the broad exposure of PHCZs
43
in San Francisco Bay and their characteristics of bioaccumulation and biomagnification
44
along with dioxin-like effects. These findings raise the need for additional research to
45
better elucidate their sources, environmental behavior, and fate in global environments.
46
ACS Paragon Plus Environment
2
Page 3 of 35
Environmental Science & Technology
47
48 49 50 51
TOC Art
52 53 54 55 56 57 58 59 60 61 62 63 64 65
ACS Paragon Plus Environment
3
Environmental Science & Technology
66
Page 4 of 35
INTRODUCTION
67
Marine and freshwater systems are subject to contamination by a large variety of
68
halogenated substances. Well-known halogenated contaminants that have been
69
extensively investigated worldwide, such as polybrominated diphenyl ethers (PBDEs)
70
and polychlorinated biphenyls (PCBs) as well as many others, have been added to the
71
persistent organic pollutants (POPs) list of the Stockholm Convention, due to their global
72
contamination and toxic potencies.1-4 Recently, a number of polyhalogenated carbazoles
73
(PHCZs) were discovered in soil and aquatic systems from different regions. Chlorinated
74
and brominated carbazoles, including 3-chlorocarbazole (3-CCZ), 3-bromocarbazole (3-
75
BCZ), 3,6-dibromocarbazole (36-BCZ), 1,3,6,8-tetrabromocarbazole (1368-CCZ) or
76
1,3,6,8-tetrabromocarbazole (1368-BCZ), were reported in sediment from the Laurentian
77
Great Lakes of North America, the Saginaw River system (United States or U.S.), the
78
Ontario River (Canada), the North Sea Estuary (Germany), the Lippe River (Germany)
79
and Lake Tai (China), as well as in soil from Germany and Greece.5-15 Carbazoles
80
substituted with a combination of chlorine and bromine atoms (e.g., 1,8-dibromo-3,6-
81
dichlorocarbazole or 18-B-36-CCZ) were also found in sediment from Southern Ontario,
82
Canada and the Great Lakes.5,6,9 The Great Lake studies identified a number of other
83
PHCZs with various combinations of halogen (bromine, chlorine, and iodine)
84
substitutions, including Br2-, Br3-, Br4-, Br3Cl-, Br3ClI-, Br4Cl-, Br4I-, and Br5-
85
carbazole.5,6 To date, the potential sources of the variety of PHCZ substances have not
86
been thoroughly studied. Halogenated indigo dyes have been suggested as potential
87
sources of 1368-BCZ and 18-B-36-CCZ, as well as some other PHCZs.16 These PHCZs
88
may be formed as impurities in the final products of halogenated indigo dye formulations.
ACS Paragon Plus Environment
4
Page 5 of 35
Environmental Science & Technology
89
Natural origins have also been suggested for selected PHCZs or their precursors.5,17
90
Selected PHCZs can be produced through enzymatic synthesis of bromo- and chloro-
91
carbazoles by chloroperoxidase from marine fungus Caldariomyces fumago in water.18
92
But the extreme scarcity in relevant information and studies made it impossible to
93
elucidate the sources of PHCZs and the volume of anthropogenic input.
94
In addition to the discovery of PHCZs in environmental compartments, recent
95
studies also investigated toxicological properties of selected halogenated carbazoles,
96
including dioxin-like effects, and developmental, cardiotoxic, and mutagenic activities.19-
97
22
98
injection dose of 30 or 60 mg/kg body weight (bw) for five consecutive days or at a
99
single injection of 50 – 300 mg/kg bw resulted in the induction of a dominant lethal
100
mutation or abnormal sperm heads, respectively.22 In zebrafish (Danio rerio), chronic
101
exposure to both 27-BCZ and 2367-CCZ resulted in developmental toxicity in embryos at
102
nano-molar levels and induced the expression of cytochrome P450 enzyme CYP1A1 in
103
the heart area at micro-molar levels.20 The aryl hydrocarbon receptor (AhR)-mediated
104
effects in MDA-MD-468 human breast cancer cells were also reported for 3-CCZ, 36-
105
CCZ, 27-BCZ, 2367-CCZ and 18-B-36-CCZ, as well as some other PHCZs.19
Treatment of adult male mice (Mus musculus) with carbazole at a daily intraperitoneal
106
Despite an increasing number of reports on the environmental occurrence and
107
toxicity of PHCZs, knowledge of their bioavailability and consequent contamination in
108
aquatic and terrestrial organisms remains extremely limited. The present work introduces
109
the first comprehensive study to address the contamination of PHCZs in abiotic and
110
biotic components of one of the most studied estuarine systems in North America, San
111
Francisco Bay (U.S.). San Francisco Bay has been the subject of long-term monitoring of
ACS Paragon Plus Environment
5
Environmental Science & Technology
Page 6 of 35
112
anthropogenic pollution, which revealed some of the highest concentrations of POPs
113
(e.g., PBDEs and PCBs) worldwide in its sediment and biota, mainly due to a dense
114
urban population and associated activities.23-27 The specific objectives of this study were
115
to: (1) investigate spatial distributions of PHCZs in San Francisco Bay sediment; (2)
116
evaluate the bioaccumulation and biomagnification potency of PHCZs; and, (3)
117
fingerprint the composition patterns of PHCZ congeners in sediment and biota.
118 119
MATERIALS AND METHODS
120
Chemicals and Reagents. The reference standards of 3-CCZ, 36-CCZ, 1368-CCZ, 2367-
121
CCZ, 1,3,6-tribromocarbazole (136-BCZ), 1-bromo-3,6-dichlorocarbazole (1-B-36-
122
CCZ), and 18-B-36-CCZ were purchased from Wellington Laboratories (Guelph, ON,
123
Canada). The standards of 3-BCZ, 27-BCZ, and 36-BCZ were purchased from Sigma-
124
Aldrich (St. Louis, Missouri). The reference standard of carbazole and 1368-BCZ were
125
purchased from AccuStandard (New Haven, CT) and Florida Center for Heterocyclic
126
Compounds of the University of Florida (Gainesville, FL), respectively. Surrogate
127
standards 4'-fluoro-2,3',4,6-tetrabromodiphenyl ether (F-BDE69), 4'-Fluoro-2,3,3',4,5,6-
128
hexabromodiphenyl ether (F-BDE160) and 2,2',3,4,4',5,6,6'-octachorobiphenyl (PCB-
129
204), as well as internal standards 3'-Fluoro-2,2',4,4',5,6'-hexabromodiphenyl ether (F-
130
BDE154) and decachlorodiphenyl ether (DCDE), were purchased from AccuStandard
131
(New Haven, CT). A total of 20 PBDE congeners, including BDE-28, -47, -49, -66, -85,
132
-99, -100, -138, -153, -154, -183, -196, -197, -201, -202, -203, -206, -207, -208 and -209,
133
were purchased from AccuStandard. Diatomaceous earth and sodium sulfate (10-60
134
mesh) were purchased from Fisher Scientific (Hanover Park, IL) and treated in a muffle
ACS Paragon Plus Environment
6
Page 7 of 35
Environmental Science & Technology
135
furnace at 600 °C overnight (> 12 h) prior to use. Copper (50 mesh, granular, reagent
136
grade), as well as high performance liquid chromatography (HPLC) grade solvents, was
137
purchased from Fisher Scientific. Isolute® silica sorbent (average pore size: 60) was
138
purchased from Biotage Inc. (Charlotte, NC, USA) and baked at 130 °C prior to use.
139
Samples. Surficial sediment was collected at a depth of 0 to 5 cm using a Young-
140
modified Van Veen grab with a surface area of 0.1 m2 at 26 different sites in San
141
Francisco Bay in 2014 (Figure 1A; Table S2).23 Sediment was kept in pre-cleaned I-
142
Chem jars after collection. A field blank was also prepared along with sediment
143
collection. Transplanted bivalves (Mytilus californianus) were collected from a non-
144
urban reference site (Tomales Bay; 38°18'00.0"N, 123°04'12.0"W) and deployed at six
145
stations in the Bay for 90 days during the summer of 2014 (Figure 1B; Table S2).
146
Approximately 30-40 bivalves were sampled from each site and their soft tissues were
147
homogenized. Four species of sport fish, including white croaker (Genyonemus lineatus),
148
striped bass (Morone saxatilis), jacksmelt (Atherinopsis californiensis), and white
149
sturgeon (Acipenser transmontanus), were collected at 8 recreational fishing areas across
150
the Bay in 2014 and 2009 (Figure 1B; Table S2). Tissues from 1 to 21 individual fish of
151
each species from the same site and collection year were homogenized to make a
152
composite. Jacksmelt were processed whole but with the head, tail, and viscera removed.
153
White croaker were processed whole, and all other samples were processed as muscle
154
fillets with skin removed. A total of 12 fish composites were processed for chemical
155
analysis. Eggs of double-crested cormorants (Phalacrocorax auritus) were collected in
156
2016 at three locations within the Bay: Wheeler Island (North Bay), Richmond Bridge
157
(Central Bay), and Don Edwards Wildlife Refuge (South Bay) (Figure 1B; Table S2).
ACS Paragon Plus Environment
7
Environmental Science & Technology
Page 8 of 35
158
Two to three egg composites, containing 7 – 10 eggs per composite, were available from
159
each site, resulting in a total of eight composites for chemical analysis. Blubber tissues
160
were sampled in 2007 – 2015 from 18 adult or sub-adult harbor seals residing in the Bay
161
(Figure 1B; Table S2). These harbor seals had various conditions during blubber
162
collection, i.e. captured alive (n=5), deceased during treatment (n=3), euthanized (n=3),
163
or deceased in the wild (n=7). Detailed blubber collection method is provided in the
164
Supporting Information. All samples were shipped with ice packs to the analytical
165
laboratory at the Southern Illinois University and stored at -20 °C prior to chemical
166
analysis.
167
Sample Preparation. Sample pretreatment and cleanup procedures described
168
below applied to the determination of both PHCZ and PBDE congeners. Sediment
169
samples were freeze-dried for 48 hours and sieved through a 100-micron stainless cloth
170
sieve (Hogentogler & Co. Inc., Columbia, MD). Dry sediments were analyzed for the
171
total organic carbon (TOC) percentages using a FlashEA® 1112 Nitrogen and Carbon
172
Analyzer (Thermo Fisher Scientific, Waltham, MA). Bivalve composites were freeze-
173
dried for 48 hours. Approximately 5 g of dried sediment, 1.0 g of bivalve and fish
174
composites, 3.0 g of egg composites, or 0.11 – 0.49 g of blubber was ground with
175
diatomaceous earth. After spiking with surrogate standards (FBDE-69, FBDE-160, and
176
PCB-204; 50 ng each), the sample was subjected to accelerated solvent extraction
177
(Dionex ASE 350, Sunnyvale, CA, USA) with dichloromethane (DCM) at 100 °C and
178
1500 psi. Copper powder activated by concentrated hydrochloric acid was used to
179
remove sulfur in the sediment extract.8 Tissue extracts were subjected to gravimetric
180
determination of lipid content by using 10% of the extract. The extract after copper
ACS Paragon Plus Environment
8
Page 9 of 35
Environmental Science & Technology
181
treatment or lipid determination was purified by a Shimadzu Prominence Semi-Prep
182
HPLC (Shimadzu America Inc., Columbia, MD) equipped with a Phenogel gel
183
permeation chromatography (GPC) column (300 × 21.2 mm, 5µ, 100Å; Phenomenex,
184
Inc., Torrance, CA). The mobile solvent was 100% DCM and the flow rate was set at 4
185
mL/min. Target compounds (including PHCZ and PBDE congeners) were collected in
186
the fraction ranging from 16 to 45 minutes. The resulting fraction was further cleaned and
187
separated on a 2-g Isolute® silica solid phase extraction (SPE) column packed into a 6 mL
188
polyesters SPE tube (SiliCycle Inc., Quebec City, Canada). The SPE column was pre-
189
washed with 10 mL hexane (HEX) to condition the silica gel sorbent. After the sample
190
was loaded, the first fraction was eluted with 3 mL HEX and was discarded. The second
191
fraction that contained target PHCZs was eluted with 11 mL of a mixture of HEX and
192
DCM (40:60, v/v). The latter fraction was concentrated to approximately 200 µL and
193
transferred to an insert tube in a gas chromatography (GC) vial. Internal standards FBDE-
194
154 and DCDE (100 ng each) were added prior to instrumental analysis.
195
Instrumental Analysis. The separation and quantification of the target PHCZs, as
196
well as the surrogate and internal standards, was performed on an Agilent 7890B gas
197
chromatography (GC; Agilent Technologies, Palo Alto, CA) coupled to a single
198
quadrupole mass analyzer (Agilent 5977A MS) in either electron-capture negative
199
ionization (ECNI) or electron impact (EI) mode. Our previous study has revealed that 3-
200
CCZ, 36-CCZ, or 3-BCZ exhibited a greater sensitivity in EI versus ECNI mode, whereas
201
the other PHCZ congeners had a much better sensitivity in ECNI mode.12 The column
202
used for GC-MS analysis was a 30 m HP-5MS column (0.25 mm i.d., 0.25 µm, J&W
203
Scientific, Agilent Tech.). The injector was operated in pulsed-splitless mode, held at
ACS Paragon Plus Environment
9
Environmental Science & Technology
Page 10 of 35
204
260 °C. Initial oven temperature was held at 50 °C for 3 min, increased to 150 °C at 10
205
°C/min, and then ramp to 300 °C at 5 °C/min and held for 10 min. The quantification and
206
confirmation of each target compound was achieved via selected ion monitoring (SIM)
207
for its characteristic ions under ECNI or EI mode (Table S1). Instrumental analysis of
208
PBDE congeners was described in detail in Supporting Information.
209
Quality Assurance and Control. The QA/QC practices in the present study
210
included the analysis of field or laboratory procedural blanks, spiking recovery tests, and
211
replicates of randomly chosen authentic samples, as well as monitoring the recoveries of
212
surrogate standards. No PHZCs were detected in any field or laboratory blanks. Analyses
213
of reference sediment (collected from Au Sable River, Michigan, USA), fish (Tilapia,
214
Oreochromis niloticus) fillets, and chicken (Gallus domesticus) eggs spiked with known
215
amounts of PHCZs (25 ng of each congener) revealed a mean recovery of individual
216
PHCZ congeners ranging from 71.6% to 113.8% (adjusted with surrogate recoveries) in
217
all matrices. Fish fillets and chicken eggs were purchased from a local supermarket and
218
were confirmed to be free of PHCZs prior to spiking experiments. No PHCZs were
219
quantifiable in reference sediment.
220
revealed the relative standard deviations (RSDs) less than 6.7% for the summed
221
concentrations of detected PHCZ congeners (referred to ΣPHCZs) and ranged from 4.8%
222
to 9.1% for individual congeners. Recoveries of surrogate standards were 92.3 ± 13.5%
223
(mean ± standard deviation) for FBDE-160 and 88.8 ± 11.4% for PCB-204, respectively.
224
The method limits of quantification (MLOQs) were assessed by multiplying a Student’s
225
t-value designated for a 99% conference level with standard deviations in the replicate
226
analyses (n = 8) of reference sediment, fish, or eggs, following the protocol introduced in
Replicate analysis of random samples (n = 3)
ACS Paragon Plus Environment
10
Page 11 of 35
Environmental Science & Technology
227
Wu et al.8 The MLOQs ranged from 0.1 to 0.3 ng/g dw in sediment, 0.8 – 1.5 ng/g lw in
228
fish, and 0.5 – 1.2 ng/g lw in bird eggs (Table S1). The MLOQs in bivalve and harbor
229
seal were adjusted from those determined in fish with relative lipid contents. Analytes
230
with instrumental responses below the instrumental detection limit (IDL), defined as a
231
concentration yielding a signal-to-noise ratio (S/N) of 5, were considered non-detectable
232
(nd). For measurements below MLOQs or non-detectable for an analyte with detection
233
frequency above 60%, a regression plotting method was applied to assign values for
234
statistical analysis.28
235
Data Analysis. The PHCZ concentrations in San Francisco samples were
236
adjusted with recoveries of surrogate standards (i.e. FBDE-160 for ECNI and PCB-204
237
for EI analyses) and reported as ng/g dry weight (dw) for sediment and ng/g lipid weight
238
(lw) for biological samples. The suitability of using these surrogate standards for
239
recovery adjustment has been demonstrated in previous studies.8,13 The Student’s t-test
240
and the Pearson’s correlative analyses, as well as the principal component analysis
241
(PCA), were conducted using the OriginPro 9.0 (OriginLab Corporation). The analysis of
242
variance (ANOVA) was conducted using the PASW Statistics 18.0 (IBM Inc.). The level
243
of significance is set as α = 0.05. Concentrations of PBDEs in all samples of the present
244
study have not been published elsewhere.
245 246
RESULTS AND DISCUSSION
247
PHCZs in San Francisco Bay Sediment. PHCZs were detected in sediment samples
248
collected from all 26 studied sites, revealing their wide distribution in the Bay (Figure
249
1A). The concentrations of ΣPHCZs ranged from 1.7 to 20.5 ng/g dw (Table S3). PHCZs
ACS Paragon Plus Environment
11
Environmental Science & Technology
Page 12 of 35
250
were also reported in other aquatic systems,5-13 but the number of investigations to date
251
remains overall very limited. The median concentration (i.e., 9.3 ng/g dw) of ΣPHCZs in
252
the Bay was lower than what has been reported in surface sediment from the Saginaw
253
River basin (Michigan, U.S.; median: 18.9 ng/g dw) and the Great Lakes (median: 38.0
254
ng/g dw), but much greater than the levels reported in sediment collected from rivers and
255
coastal water of the North Sea estuary (Germany; median: 0.23 ng/g dw) and Lake Tai of
256
China (median: 1.5 ng/g dw).5,8,10,13 It should be noted that the Great Lakes study
257
included a number of additional congeners compared to those determined in the present
258
study, which do not have corresponding commercial reference standards available and
259
were subjected to semi-quantification.5 Selected PHCZ congeners were also identified,
260
quantified, or semi-quantified in sediment from the Ontario River (Canada) and Lake
261
Michigan (U.S.) sediment cores.6,7,9 Their detection in various watersheds may suggest a
262
broad distribution of these chemicals in global aquatic ecosystems.
263
Spatial distribution of PHCZs revealed some elevated concentrations at sites
264
along the eastern side of the Central and South Bay, agreeing with previous findings for
265
PBDEs and PCBs at these sites.23,24 These elevated concentrations were likely due to
266
greater urbanization as well as geographic and hydrological features. The South Bay
267
extends from the Central Bay toward San Jose, while the North Bay extends from the
268
freshwaters at the mouths of the Sacramento and San Joaquin rivers through the saline
269
waters of Suisun and San Pablo Bays.25 Central and South Bay waterfronts include the
270
majority of the industrial, port, and shipping activities in the region.26 Highly populated
271
metropolitan centers including San Francisco and Oakland are located in the Central Bay,
272
while San Jose is located in the South Bay. The Central and South Bay segments in total
ACS Paragon Plus Environment
12
Page 13 of 35
Environmental Science & Technology
273
receive approximately 76% of the region’s wastewater inflow into the Bay.25,27 Given
274
that some PHCZs were suggested as impurities in the final products of historically
275
manufactured halogenated indigo dye formulations,16 they may be present in industrial or
276
municipal wastewater. Hydrological features may also be important. Surface waters in
277
the South Bay experience the least amount of mixing with non-wastewater effluent flow,
278
particularly in the dry seasons, and thus have higher hydraulic residence time relative to
279
other Bay segments.26 By contrast, the northern Bay segment receives up to 90% of the
280
Estuary’s freshwater inflow;25 while these river discharges are influenced by upstream
281
urban contaminants, they may nevertheless dilute environmental contaminants
282
concentrated in local urban runoff or wastewater effluent and result in short residence
283
time.
284
However, statistical analysis did not reveal a significant difference in dry weight
285
(p = 0.13) or TOC (p = 0.35) based ΣPHCZ concentrations among the North, Central, and
286
South Bay segments, although a north-to-south increasing gradient has been reported for
287
BDE-209 and PCBs.23,24 The TOC of sediment significantly correlated with ΣPHCZ
288
concentrations (p < 0.001), but did not differ significantly among the Bay segments. Of
289
note, a few sites from the northern segment exhibited sediment concentrations above the
290
median value (Figure 1A). Some of the northern sites were also reported to have elevated
291
PBDE concentrations in sediment.23 Possible anthropogenic sources or natural origins
292
may be present in the North Bay, but current information is insufficient to draw any
293
conclusion.
294 295
PHCZs Are Bioavailable and Biomagnify.
PHCZs were detected in all
biological samples, including transplanted bivalves, sport fish, cormorant eggs, and
ACS Paragon Plus Environment
13
Environmental Science & Technology
Page 14 of 35
296
harbor seal blubber (Figure 2; Table S3). Bivalves such as mussels are stationary
297
organisms that filter particles from water. Contaminant concentrations in bivalve tissues
298
are not only a good indicator of local contamination, but also an ideal measurement of the
299
bioavailability of environmental pollutants because bivalves generally perform little
300
metabolic transformation.29 Concentrations of ΣPHCZs in bivalve composites from the
301
six sites within the Bay ranged from 8.3 to 76.1 ng/g lw, with a median concentration of
302
33.7 ng/g lw. These levels were greater than that in the reference site (6.8 ng/g lw)
303
located in the Tomales Bay, indicating elevated contamination in the estuarine system
304
subject to urban influences. It is also noted that bivalves studied in the present work were
305
transplanted to the designated sites, thus subject to shorter exposure time than resident
306
bivalves, which may be exposed at a greater level.
307
PHCZs were also present in sport fish, cormorant eggs, and harbor seal blubber, at
308
median concentrations of 53, 155, and 164 ng/g lw, respectively. The occurrence in
309
various species occupying different trophic levels, from benthic invertebrates to top
310
predators, demonstrates substantial bioavailability of PHCZs and their broad exposure in
311
the Bay ecosystem. The estimated bioconcentration factors (BCFs) of PHCZs via the
312
U.S. Environmental Protection Agency (EPA) Estimation Program Interface (EPI) Suite
313
Version 4.11 have logarithmically transformed values ranging from 2.73 to 3.15 (Table
314
S1), except for 3-CCZ (2.43) and 3-BCZ (2.47). These estimated values are below the
315
criteria for bioaccumulation (i.e. log BCF = 3.3) recommended by the European
316
Commission Registration, Evaluation and Authorization of Chemicals (REACH)
317
program.30 However, studies have suggested that a conservative interval of 0.75 log units
318
should be considered when employing the BCF criteria, given the variability of the
ACS Paragon Plus Environment
14
Page 15 of 35
Environmental Science & Technology
319
experimental BCF data used for model predictions.31,32 Thus, a chemical with estimated
320
log BCF less than 3.3 cannot be safely classified as not bioaccumulative. Indeed, the
321
Arnot/Gobas bioaccumulation factor (BAF) model suggests that an organic chemical with
322
log Kow (octanol-water partitioning coefficient) value > 4 may possess bioaccumulation
323
potential in aquatic food webs.30 Given the lack of experimentally determined data, the
324
EPA EPI program was used to calculate log Kow for the analyzed PHCZ congeners, which
325
range from 4.12 to 6.85, except for 3-CCZ (calculated log Kow = 3.94). Therefore, both
326
model predictions and our environmental findings suggest considerable bioavailability
327
for the majority, if not all, of PHCZ congeners.
328
The data suggest biomagnification of PHCZs from fish to harbor seal. Three of
329
the four investigated sport fish species, including white croaker, striped bass, and
330
jacksmelt, have been demonstrated to constitute a portion of the diet of Pacific harbor
331
seals which are opportunistic feeders, even though they do not represent seals’ major
332
food sources.33 The biomagnification factor (BMF) of ΣPHCZs, calculated as the ratio of
333
median ΣPHCZ concentration in seal blubber to that in fish composites of each species,
334
was determined to be 6.1, 3.6 and 2.7 from white croaker, jacksmelt or striped bass to
335
seal, respectively. The BMFs for ΣPHCZs were even greater than those of ΣPBDEs in the
336
same fish-seal food chains (i.e., 1.0 – 2.9). The BMFs were also determined for
337
individual PHCZ congeners (Table S4). The results indicated that the biomagnification of
338
PHCZs was mainly driven by chlorinated carbazoles, particularly 36-CCZ which
339
exhibited statistically greater BMFs (3.3 – 7.5) than other congeners (i.e., 1368-CCZ, 36-
340
BBZ and 136-BCZ; ANOVA with Fisher’s post-hoc test, F3,8 = 11.5, p = 0.003). The
341
marine mammal food web model developed by Kelly et al. suggests that
ACS Paragon Plus Environment
15
Environmental Science & Technology
Page 16 of 35
342
biomagnification capacities of organic chemicals are primarily controlled by their
343
physicochemical properties, such as log Kow and log Koa, assuming no metabolic
344
transformation.34 The chemicals that are subject to great levels of biomagnification in
345
marine mammalian food webs normally have a log Kow ranging from ~4 to ~7.8 and a log
346
Koa > 7.34
347
biomagnification potential in marine mammalian food webs. However, most of the BMF
348
values determined for brominated carbazoles were less than one in our study, likely
349
suggesting elevated metabolism of these congeners in harbor seals. But additional studies
350
are required to support this hypothesis. We would also like to point out that the
351
abovementioned approach used to determine BMFs has various sources of uncertainty
352
and may be affected by a number of factors,35 including the non-dominance of the
353
investigated fish species in seals’ diet, variations in age, sex, or collection year of the
354
analyzed seal samples or variations in the species and collection year of fish composites,
355
as well as relatively small sample sizes for each species. These sources of uncertainty
356
may account for a large variability of the measured BMF values between studies. Indeed,
357
the measured BMFs for ΣPBDEs in the present study (1.0 – 2.9) were generally lower
358
than the previously reported values (i.e. 2 – 76) in fish – seal food chains from the
359
northwestern Atlantic, North Sea, and Svalbard (Norway).36-41 Nevertheless, a direct
360
comparison of the BMF values between PHCZs and PBDEs in the same food chains in
361
the present study may adequately demonstrate the biomagnification potency of PHCZs,
362
mostly driven by chlorinated congeners, in the studied ecosystem. Future studies are
363
needed to better elucidate the biomagnification or metabolism of individual PHCZ
364
congeners through prey-predator food chains or different trophic levels of a food web.
Most PHCZs meet these criteria and therefore may possess substantial
ACS Paragon Plus Environment
16
Page 17 of 35
365
Environmental Science & Technology
Congener-Specific
Bioaccumulation
in
Different
Species.
Congener
366
compositions of PHCZs differed among studied species and sediment (Figure 3). The
367
PCA analysis of congener compositions revealed three clusters in the biplot (Figure 4).
368
Congener compositions of bivalves resembled closely that of sediment (Cluster I), which
369
were dominated by 36-CCZ, followed by 36-BCZ, but had elevated contributions by
370
1368-BCZ compared to other species. Fish and harbor seal (Cluster II) shared a similar
371
congener composition profile, which exhibits a dominance of 36-CCZ, followed by 136-
372
BCZ and then 36-BCZ. Cormorant eggs (Cluster III) differed from all other species in
373
composition profiles, and were dominated by 136-BCZ, followed by 36-CCZ and 36-
374
BCZ. The elevated composition of 136-BCZ may indicate its greater biomagnification
375
potential in piscivorous bird food chains compared to that in marine mammal food chains
376
or greater metabolism by harbor seal than cormorant. The differences revealed by PCA
377
may overall suggest congener-specific bioaccumulation, biomagnification, or metabolism
378
among benthic species (Cluster I), fish and mammals (Cluster II), and avian species
379
(Cluster III) in Bay food webs.
380
The dominance of 36-CCZ in most species and sediment from the Bay was
381
consistent with previous findings in sediment from some other regions.8,10,13 This
382
consistency may indicate a relative abundance of this specific PHCZ congener during the
383
synthesis and manufacturing or its stability during degradation. These characteristics may
384
be attributed to an electrophilic aromatic substitution pattern which favors halogen
385
substitution at the ortho and para positions relative to nitrogen on the carbazole, as well
386
as the enhanced stability of para-substituted products resulting from the high charge
387
density at the para positions compared to the ortho positions and the effect of the lone
ACS Paragon Plus Environment
17
Environmental Science & Technology
Page 18 of 35
388
pair of electrons in the NH2 moiety.18,42,43 Indeed, a dominance of products with halide
389
substitution at the 3,6- positions has been found in enzymatic synthesis of PHCZs from a
390
source material of carbazole.18 The 36-BCZ was also frequently detected in the Bay, and
391
was the second dominant congener in bivalves and sediment.
392
Although cormorant eggs and harbor seal blubber revealed comparable
393
concentrations of ΣPHCZs, their composition profiles differed significantly, mainly due
394
to a dominance of 136-BCZ over other congeners in the cormorants. Although additional
395
evidence is needed, our data suggest congener-specific bioaccumulation/biomagnification
396
potential or metabolic differences between marine mammal and avian food webs in the
397
studied system. The effect of depuration kinetics (e.g., maternal transfer in mammals and
398
birds) on composition profiles cannot be ignored, but it needs further investigations.
399
Unfortunately, comparison with other studies regarding the accumulation pattern of
400
PHCZs is not yet possible due to the extreme scarcity of relevant knowledge. While our
401
work provides insights in the bioaccumulation of PHCZs in aquatic organisms, additional
402
efforts are essential to better elucidate the trophodynamics and fate of PHCZs in different
403
biological systems.
404
Implications and Future Research Needs. To better understand the abundance
405
of PHCZs relative to other important anthropogenic pollutants in San Francisco Bay, we
406
compared their concentrations in sediment and biological tissues with three groups of
407
POP substances: PBDEs, PCBs and hexachlorohexanes (HCHs). These contaminants
408
have been broadly distributed in global ecosystems and subjected to long-term
409
environmental monitoring in many important aquatic ecosystems, including San
410
Francisco Bay. Concentrations of ΣPHCZs were significantly greater (p < 0.001) than
ACS Paragon Plus Environment
18
Page 19 of 35
Environmental Science & Technology
411
that of ∑PBDEs (summed concentration of detected PBDE congeners) in the same
412
sediment samples (0.2 – 5.8 ng/g dw) (Figure S1, Supporting Information). The median
413
concentration ratio of ΣPHCZs to ∑PBDEs was 5.2 in sediment, while the ratios ranged
414
from 0.10 to 0.29 in organisms examined in the present study. Significant correlations
415
(Figure 5) were also found between ∑PHCZ and ∑PBDE concentrations in the sediment
416
and biota (p < 0.01 in all species except for transplanted bivalves), suggesting that these
417
two groups of contaminants may share some common sources (e.g., wastewater effluent
418
and urban runoff) and possess similar environmental processes.
419
Concentrations of PHCZs were greater than those of ΣHCHs and comparable to
420
those of ΣPCBs in the sediment of San Francisco Bay (Figure S1). 44,45 Data of sediment-
421
associated PCBs and HCHs included in Figure S1 were from multiple sites, even though
422
the sites were not entirely consistent with those analyzed in the present study. The levels
423
of PHCZs were also greater than those of ΣHCHs in various biological species from the
424
Bay ecosystem, but were generally two to three orders of magnitude lower than ΣPCBs in
425
biota.44-46 It is noted that the number of PCB congeners under investigation differed
426
among studies, ranging from 46 to 209.44-46 The median concentration ratios of ΣPHCZs
427
to ∑PCBs and ΣPHCZs to ∑HCHs in the Bay organisms ranged from 0.004 to 0.07 and
428
1.7 – 10.9, respectively.44-46 These comparisons suggest PHCZs are among the groups of
429
halogenated contaminants with considerable abundances in the Bay aquatic ecosystems.
430
PBDE concentrations in the Bay have declined by approximately half in sport fish and
431
74-95% in bivalves and bird eggs, as a result of discontinuation in production since
432
2004.23 Concentrations of other legacy POPs (e.g., organochlorine pesticides) have also
433
declined over time in the Bay system.47 The future trends of PHCZ concentrations in Bay
ACS Paragon Plus Environment
19
Environmental Science & Technology
Page 20 of 35
434
sediment and biota remain unknown due to the lack of temporal studies, thus warranting
435
further investigation.
436
Among a variety of potential toxicity activities of PHCZs, their dioxin-like
437
activity has raised particular concern.19-21,48 Riddell et al. reported a similarity between
438
PHCZs and other dioxin-like compounds in the structure-dependent induction of
439
cytochrome P450 (CYP450) 1A1 and CYP1B1 gene expression in aryl hydrocarbon-
440
responsive MDA-MB-468 breast cancer cells and that the magnitude of the induction
441
responses for the most active PHCZs (e.g., 2367-CCZ) was reported to be similar to that
442
for 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD).19 The relative effect potencies (REP) of
443
PHCZs (compared to TCDD) were estimated to range from 0.000013 to 0.00066 for
444
mono- to tetra-PHCZs.19 The REP can be used to represent toxic equivalency factors
445
(TEFs) of PHCZs, given that they were determined following an extensively used
446
approach for TEF estimations.49 Based on the REP measurements and using the equation
447
described in Wu et al. (also seen in Supporting Information),13 we estimated the median
448
toxic equivalent (TEQ) of PHCZs to be 1.2 pg TEQ/g dw in sediment and 4.8 – 19.5 pg
449
TEQ/g lw in biological tissues (Table S3). The TEQs of PHCZs appeared to be lower
450
than those reported for polychlorinated dibenzo-dioxins/furans (PCDDs/Fs) and co-planar
451
PCBs (i.e. 33.1 and 109 pg/g lw, respectively) in fish from California coastal waters
452
(collected in 2000-2001).50 Using the same approach, Wu et al. has estimated the TEQs
453
of PHCZs in Lake Tai (China) sediments ranging from 0 to 1.36 pg TEQ/g dw (median =
454
0.12).13 Pena-Abauurea et al. reported the TEQ for 18-B-36-CCZ alone ranging from
455
0.31 to 7.46 pg TEQ/g dw in stream sediment from Ontario (Canada),48 by employing a
456
REP of 0.0016 for this particular congener, different from what we used (0.0003) in the
ACS Paragon Plus Environment
20
Page 21 of 35
Environmental Science & Technology
457
present study. Future studies are needed to unify the TEF values for the evaluation of
458
dioxin-like effects of PHCZs in global ecosystems.
459
In summary, this study has found that PHCZs are abundant in sediment of the San
460
Francisco Bay, and furthermore, they appear to have the ability to bioaccumulate and
461
biomagnify in marine mammal food webs. Bioaccumulation and biomagnification of
462
PHCZs may also be likely in terrestrial mammalian and human food webs according to
463
their physicochemical properties. These characteristics and existing evidence suggest the
464
PHCZs as a potential group of persistent organic pollutants with dioxin-like potency and
465
global distribution. However, knowledge gaps remain in many aspects of our current
466
understanding of these substances, including, but not limited to: (1) potential sources,
467
origins, or fate in aquatic systems; (2) occurrence and bioaccumulation in other
468
marine/freshwater ecosystems, non-aquatic systems and human living environments; (3)
469
rates of environmental degradation and metabolism; (4) long range transport potential;
470
and (5) other toxicities in addition to what has been reported. Thus, additional field and
471
laboratory-based research is critically needed in order to comprehensively assess the
472
environmental distribution, behavior, fate and effects of PHCZs in abiotic and biological
473
systems.
474 475
ACKNOWLEDGEMENT
476
The authors would like to acknowledge the San Francisco Estuary Institute and the
477
Regional Monitoring Program for Water Quality in San Francisco Bay (RMP) for
478
samples and financial support. We thank J. Davis, J. Sun, P. Trowbridge, and D. Yee for
479
assistance and constructive comments. We also thank D. Greig and B. Halaska for
ACS Paragon Plus Environment
21
Environmental Science & Technology
Page 22 of 35
480
providing archived harbor seal blubber samples collected by The Marine Mammal Center
481
and California Academy of Sciences under stranding agreements from The Marine
482
Mammal Health and Response Program and J. Harvey, Moss Landing Marine
483
Laboratories, for wild-caught harbor seal samples collected under NMFS research permit
484
#16991. A. Rothert and S. Baer from SIU Core Facility for Ecological Analyses were
485
thanked for TOC analysis.
486 487
Supporting Information
488
Table S1-S4 and Figure S1 are available in the Supporting Information.
489 490
REFERENCES
491
(1) de Wit, C.A. An overview of brominated flame retardants in the environment.
492
Chemosphere 2002, 46, 583-624.
493
(2) Law, R.J.; Herzke, D.; Harrad, S.; Morris, S.; Bersuder, P.; Allchin, C.R. Levels and
494
trends of HBCD and BDEs in the European and Asian environments with some
495
information for other BFRs. Chemosphere 2008, 73, 223-241.
496
(3) United Nations. Revised draft guidance for the inventory of polychlorinated diphenyl
497
ethers under the Stockholm Convention. Conference of the Parties to the Stockholm
498
Convention on Persistent Organic Pollutants Seventh Meeting, Geneva, 4-15 May
499
2015.
500
http://chm.pops.int/Implementation/NIPs/Guidance/GuidancefortheinventoryofPBDE
501
s/tabid/3171/Default.aspx (accessed October 2016).
Available
ACS Paragon Plus Environment
at:
22
Page 23 of 35
Environmental Science & Technology
502
(4) Ritter, L.; Solomon, K.R.; Forget, J.; Stemeroff, M.; O’Leary, C. Persistent Organic
503
Pollutants. An assessment report on: DDT Aldrin Dieldrin Endrin Chlordane
504
Heptachlor Hexachlorobenzene Mirex Toxaphene Polychlorinated Biphenyls Dioxins
505
and Furan. The International Programme on Chemical Safety (IPCS) within the
506
framework of the Inter-Organization Programme for the Sound Management of
507
Chemicals
508
http://chm.pops.int/Implementation/PCBs/DocumentsPublications/tabid/665/Default.
509
aspx (accessed October 2016).
(IOMC).
Available
at:
510
(5) Guo, J.; Li, Z.; Ranasinghe, P.; Bonina, S.; Hosseini, S.; Corcoran, M.B.; Smalley,
511
C.; Rockne, K.J.; Sturchio, N.C.; Giesy, J.P.; Li, A. Spatial and temporal trends of
512
polyhalogenated carbazoles in sediments of upper Great Lakes: Insights into their
513
origin. Environ. Sci. Technol. 2016. DOI: 10.1021/acs.est.6b06128.
514
(6) Guo, J.; Chen, D.; Potter, D.; Rockne, K.J.; Sturchio, N.C.; Giesy, J.P.; Li, A.
515
Polyhalogenated carbazoles in sediments of Lake Michigan: a new discovery.
516
Environ. Sci. Technol. 2014, 48, 12807-12815.
517 518
(7) Zhu, L.; Hites, R.A. Identification of brominated carbazoles in sediment cores from Lake Michigan. Environ. Sci. Technol. 2005, 39, 9446-9451.
519
(8) Wu, Y.; Moore, J.; Guo, J.; Li, A.; Grasman, K.; Choy, S.; Chen, D. Multi-residue
520
determination of polyhalogenated carbazoles in aquatic sediments. J. Chromatogr. A
521
2016, 1434, 111-118.
522
(9) Pena-Abaurrea, M.; Jobs, K.J.; Ruffolo, R.; Shen, L.; McCrindle, R.; Helm, P.A.;
523
Reiner, E.J. Identification of potential novel bioaccumulative and persistent
ACS Paragon Plus Environment
23
Environmental Science & Technology
Page 24 of 35
524
chemicals in sediments from Ontario (Canada) using scripting approaches with
525
GC×GC-TOF MS analysis. Environ. Sci. Technol. 2014, 48, 9591-9599.
526
(10) Chen, W.L.; Xie, Z.; Wolschke, H.; Gandrass, J.; Kotke, D.; Winkelmann, M.;
527
Ebinghaus, R. Quantitative determination of ultra-trace carbazoles in sediments in the
528
coastal environment. Chemosphere 2016, 150, 586-595.
529
(11) Heim, S.; Schwarzbauer, J.; Kronimus, A.; Littke, R.; Woda, C.; Mangini, A.
530
Geochronology of anthropogenic pollutants in riparian wetland sediments of the
531
Lippe River (Germany). Org. Geochem. 2004, 35, 1409-1425.
532
(12) Kronimus, A.; Schwarzbauer, J.; Dsikowitzky, L.; Heim, S.; Littke, R.
533
Anthropogenic organic contaminants in sediments of the Lippe river, Germany.
534
Water Res. 2004, 38, 3473-3484.
535
(13) Wu, Y.; Qiu, Y.; Tan, H.; Chen, D. Polyhalogenated carbazoles in sediments from
536
Lake Tai (China): distribution, congener composition, and toxic equivalent
537
evaluation. Environ. Pollut. 2017, 220 (Pt A), 142-149.
538
(14) Grigoriadou, A.; Schwarzbauer, J. Non-target screening of organic contaminants in
539
sediments from the industrial coastal area of Kavala City (NE Greece). Water Air
540
Soil Pollut. 2011, 214, 623-643.
541
(15) Trobs, L.; Henkelmann, B.; Lenoir, D.; Reischl, A.; Schramm, K.W. Degradative
542
fate of 3-chlorocarbazole and 3,6-dichlorocarbazole in soil. Environ. Sci. Pollut. Res.
543
Int. 2011, 18, 547-555.
544
(16) Parette, R.; McCrindle, R.; McMahon, K.S.; Pena-Abaurrea, M.; Reiner, E.; Chittim,
545
B.; Riddell, N.; Voss, G.; Dorman, F.L.; Pearson, W.N. Halogenated indigo dyes: A
ACS Paragon Plus Environment
24
Page 25 of 35
Environmental Science & Technology
546
likely source of 1,3,6,8-tetrabromocarbazole and some other halogenated carbazoles
547
in the environment. Chemosphere 2015, 127, 18-26.
548 549
(17) Teuten, E.L.; Reddy, C.M. Halogenated organic compounds in archived whale oil: A pre-industrial record. Environ. Pollut. 2007, 145, 668-671.
550
(18) Mumbo, J.; Lenoir, D.; Henkelmann, B.; Schramm, K.W. Enzymatic synthesis of
551
bromo- and chlorocarbazoles and elucidation of their structures by molecular
552
modeling. Environ. Sci. Pollut. Res. Int. 2013, 20, 8996-9005.
553
(19) Riddell, N.; Jin, U.H.; Safe, S.; Cheng, Y.; Chittim, B.; Konstantinov, A.; Parette,
554
R.; Pena-Abaurrea, M.; Reiner, E.J.; Poirier, D.; Stefanac, T.; McAlees, A.J.;
555
McCrindle, R. Characterization and biological potency of mono- to tetra-halogenated
556
carbazoles. Environ. Sci. Technol. 2015, 49, 10658-10666.
557
(20) Fang, M.; Guo, J.; Chen, D.; Li, A.; Hinton, D.E.; Dong, W. Halogenated carbazoles
558
induce cardiotoxicity in developing zebrafish embryos (Danio rerio). Environ.
559
Toxicol. Chem. 2016, 35, 2523-2529.
560
(21) Mumbo, J.; Henkelmann, B.; Abdelaziz, A.; Pfister, G.; Nguyen, N.; Schroll, R.;
561
Munch, J.C.; Schramm, K.W. Persistence and dioxin-like toxicity of carbazole and
562
chlorocarbazoles in soil. Environ. Sci. Pollut. Res. Int. 2015, 22, 1344-1356.
563 564
(22) Jha, A.M.; Bharti, M.K. Mutagenic profiles of carbazole in the male germ cells of Swiss albino mice. Mutat. Res. 2002, 500, 97-101.
565
(23) Sutton, R.; Sedlak, M.D.; Yee, D.; Davis, J.A.; Crane, D.; Grace, R.; Arsem, N.
566
Declines in polybrominated diphenyl ether contamination of San Francisco Bay
567
following production phase-outs and bans. Environ. Sci. Technol. 2015, 49, 777-784.
ACS Paragon Plus Environment
25
Environmental Science & Technology
568 569
Page 26 of 35
(24) Davis, J.A.; Hetzel, F.; Oram, J.J.; McKee, L.J. Polychlorinated biphenyls (PCBs) in San Francisco Bay. Environ. Res. 2007, 105, 67-86.
570
(25) Squire, S.; Scelfo, G.M.; Revenaugh, J.; Flegal, A.R. Decadal trends of silver and
571
lead contamination in San Francisco Bay surface waters. Environ. Sci. Technol. 2002,
572
36, 2379-2386.
573
(26) Klosterhaus, S.L.; Stapleton, H.M.; La Guardia, M.J.; Greig, D.J. Brominated and
574
chlorinated flame retardants in San Francisco Bay sediments and wildlife. Environ.
575
Int. 2012, 47, 56-65.
576
(27) Oros, D.R.; Hoover, D.; Rodigari, F.; Crane, D.; Sericano, J. Levels and distribution
577
of polybrominated diphenyl ethers in water, surface sediments, and bivalves from the
578
San Francisco Estuary. Environ. Sci. Technol. 2005, 39, 33-41.
579 580
(28) Newman, M.C. Quantitative methods in aquatic ecotoxicology. Lewis Publishers, Inc., Boca Raton, pp. 21-40. 1995.
581
(29) Sericano J.L. The American oyster (Crassostrea virginica) as a bioindicator of trace
582
organic contamination. Doctoral dissertation. Texax A&M University, College
583
Station, TX. 242 pp. 1993.
584
(30) Arnot, J.A.; Gobas, F.A.P.C. A generic QSAR for assessing the bioaccumulation
585
potential of organic chemicals in aquatic food webs. QSAR Comb. Sci. 2003, 22, 337-
586
345.
587
(31) Dimitrov, S.; Dimitrova, N.; Parkerton, T.; Comber, M.; Bonnell, M.; Mekenyan, O.
588
Base-line model for identifying the bioaccumulation potential of chemicals. SAR
589
QSAR Environ. Res. 2005, 16, 531-554.
ACS Paragon Plus Environment
26
Page 27 of 35
Environmental Science & Technology
590
(32) Lombardo, A.; Roncaglioni, A.; Boriani, E.; Milan, C.; Benfenati, E. Assessment
591
and validation of the CAESAR predictive model for bioconcentration factor (BCF) in
592
fish. Chem. Cent. J. 2010, 4(Suppl 1):S1.
593
(33) Gibble, C.M.; Harvey, J.T. Food habits of harbor seals (Phoca vitulina richardii) as
594
an indicator of invasive species in San Francisco Bay, California. Mar. Mam. Sci.
595
2015, 31, 1014-1034.
596
(34) Kelly, B.C.; Ikonomou, M.G.; Blair, J.D.; Morin, A.E.; Gobas, F.A.P.C. Food web-
597
specific biomagnification of persistent organic pollutants. Science 2007, 317, 236-
598
239.
599
(35) Franklin, J. How reliable are field-derived biomagnification factors and trophic
600
magnification factors as indicators of bioaccumulation potential? Conclusions from a
601
case study on per- and polyfluoroalkyl substances. Integr. Enviro. Assess. Manage.
602
2016, 12, 6-20.
603
(36) Shaw, S.D.; Berger, M.L.; Brenner, D.; Kannan, K.; Lohmann, N.; Päpke, O.
604
Bioaccumulation of polybrominated diphenyl ethers and hexabromocyclododecane in
605
the northwest Atlantic marine food web. Sci. Total Environ. 2009, 407, 3323-3329.
606
(37) Shaw, S.D.; Berger, M.L.; Weijs, L.; Covaci, A. Tissue-specific accumulation of
607
polybrominated
608
hexabromocyclododecanes (HBCDs) in harbor seals from the northwest Atlantic.
609
Environ. Int. 2012, 44, 1-6.
diphenyl
ethers
(PBDEs)
including
Deca-BDE
and
610
(38) Boon, J.P.; Lewis, W.E.; Tjoen-A-Choy, M.R.; Allchin, C.R.; Law, R.J.; de Boer, J.;
611
Ten Hallers-Tjabbes, C.C.; Zegers, B.N. Levels of polybrominated diphenyl ether
ACS Paragon Plus Environment
27
Environmental Science & Technology
Page 28 of 35
612
(PBDE) flame retardants in animals representing different trophic levels of the North
613
Sea food web. Environ. Sci. Technol. 2002, 36, 4025-4032.
614
(39) Jenssen, B.M.; Sørmo, E.G.; Bæk, K.; Bytingsvik, J.; Gaustad, H.; Ruus, A.; Skaare,
615
J.U. Brominated flame retardants in north-east Atlantic marine ecosystems. Environ.
616
Health Perspect. 2007, 115, 35-41.
617
(40) Sørmo, E.G.; Salmer, M.P.; Jenssen, B.M.; Hop, H.; Bæk, K.; Kovacs, K.M.;
618
Lydersen, C.; Falk-Petersen, S.; Gabrielsen, G.W.; Lie, E.; Skaare, J.U.
619
Biomagnification of polybrominated diphenyl ether and hexabromocyclododecane
620
flame retardants in the polar bear food chain in Svalbard, Norway. Environ. Toxicol.
621
Chem. 2006, 25, 2502-2511.
622
(41) Weijs, L.; Dirtu, A.C.; Das, K.; Gheorghe, A.; Reijnders, P.J.H.; Neels, H.; Blust,
623
R.; Covaci, A. Inter-species differences for polychlorinated biphenyls and
624
polybrominated diphenyl ethers in marine top predators from the Southern North
625
Sea: Part 2. Biomagnificaiton in harbour seals and harbour porpoises. Environ.
626
Pollut. 2009, 157, 445-451.
627
(42) Bonesi, S.M.; Erra-Balsells, R. On the synthesis and isolation of chlorocarbazoles
628
obtained by chlorination of carbazoles. J. Heterocycl. Chem. 1997, 34, 877-889.
629
(43) Effenberger, F.; Maier, A.H. Changing the ortho/para ratio in aromatic acylation
630
reactions by changing reaction conditions: A mechanistic explanation from kinetic
631
measurements1. J. Am. Chem. Soc. 2001, 123, 3429-3433.
632 633
(44) San Francisco Estuary Institute. Contaminant Data Display & Download. Available at: http://cd3.sfei.org/ (accessed September 20, 2016).
ACS Paragon Plus Environment
28
Page 29 of 35
Environmental Science & Technology
634
(45) SFEI. 2013-2014 Annual Monitoring Results. The Regional Monitoring Program for
635
Water Quality in San Francisco Bay (RMP). Contribution #758. San Francisco
636
Estuary Institute (Richmond, CA). 2015.
637
(46) Greig, D.J.; Ylitalo, G.M.; Wheeler, E.A.; Boyd, D.; Gulland, F.M.D.; Yanagida,
638
G.K.; Harvey, J.T.; Hall, A. J. Geography and stage of development affect persistent
639
organic pollutants in stranded and wild-caught harbor seal pups from central
640
California. Sci. Total Environ. 2011, 409, 3537-3547.
641
(47) Connor, M.S.; Davis, J.A.; Leatherbarrow, J.; Greenfield, B.K.; Gunther, A.; Hardin,
642
D.; Mumley, T.; Oram, J.J.; Werme, C. The slow recovery of San Francisco Bay
643
from the legacy of organochlorine pesticides. Environ. Res. 2007, 105, 87-100.
644
(48) Pena-Abaurrea, M.; Robson, M.; Chaudhuri, S.; Riddell, N.; McCrindle, R.; Chittim,
645
B.; Parette, R.; Jin, U.-H.; Safe, S.; Poirier, D.; Ruffolo, R.; Dyer, R.; Fletcher, R.;
646
Helm, P.A.; Reiner, E.J. Environmental levels and toxicological potencies of a novel
647
mixed halogenated carbazole. Emerg. Contam. 2016, 2, 166-172.
648
(49) Van den Berg, M.; Birnbaum, L.; Bosveld, A.T.; Brunstrom, B.; Cook, P.; Feeley,
649
M.; Giesy, J.P.; Hanberg, A.; Hasegawa, R.; Kennedy, S.W.; Kubiak, T.; Larsen,
650
J.C.; van Leeuwen, F.X.; Liem, A.K.; Nolt, C.; Peterson, R.E.; Poellinger, L.; Safe,
651
S.; Schrenk, D.; Tillitt, D.; Tysklind, M.; Younes, M.; Waern, F.; Zacharewski, T.
652
Toxic equivalency factors (TEFs) for PCBs, PCDDs, PCDFs for humans and
653
wildlife. Environ. Health Perspect. 1998, 106, 775-792.
654
(50) Brown, F.R.; Winkler, J.; Visita, P.; Dhaliwal, J.; Petreas, M. Levels of PBDEs,
655
PCDDs, PCDFs, and coplanar PCBs in edible fish from California coastal waters.
656
Chemosphere 2006, 64, 276-28.
ACS Paragon Plus Environment
29
Environmental Science & Technology
Page 30 of 35
657
FIGURE LEGENDS
658
FIGURE 1. Sampling sites of sediment (A) and biological samples (B), including
659
bivalve, sport fish, cormorant egg and harbor seal, in San Francisco Bay. The left bubble-
660
map (A) shows the concentrations (ng/g dry weight) of ∑PHCZs in sediment at each site.
661
FIGURE 2. Concentrations of ∑PHCZs in sediment (ng/g dry weight) and biological
662
samples (ng/g lipid weight) from San Francisco Bay. The concentration of the bivalve
663
composite from a reference site was excluded from the graph.
664
FIGURE 3. Congener compositions of PHCZs in sediment and biological samples. Error
665
bars represent standard deviations.
666
FIGURE 4. Biplot from the principle component analysis of PHCZ congener distribution
667
patters. S, B, E, F and H represent sediment, bivalve, cormorant egg, sport fish, and
668
harbor seal, respectively.
669
FIGURE 5. Correlations between concentrations of ∑PHCZs and ∑PBDEs in sediment (r
670
= 0.41, p = 0.021), sport fish (r = 0.95, p < 0.0001), cormorant egg (r = 0.94, p < 0.001),
671
and harbor seal blubber (r = 0.71, p < 0.0001).
672 673 674
ACS Paragon Plus Environment
30
Page 31 of 35
Environmental Science & Technology
675
676 677
FIGURE 1
678 679 680
ACS Paragon Plus Environment
31
Environmental Science & Technology
Page 32 of 35
681
682 683
FIGURE 2
684 685 686 687 688 689 690 691
ACS Paragon Plus Environment
32
Page 33 of 35
Environmental Science & Technology
692 693
FIGURE 3
694 695 696 697 698 699 700 701
ACS Paragon Plus Environment
33
Environmental Science & Technology
Page 34 of 35
702 703
FIGURE 4
704 705 706 707 708 709 710 711 712
ACS Paragon Plus Environment
34
Page 35 of 35
Environmental Science & Technology
713 714
FIGURE 5
ACS Paragon Plus Environment
35